The present invention relates to mitigating the toxic effect of inorganic contaminants in contaminated sites, and more particularly, to using stabilized zero-valent iron nanoparticles for the in situ immobilization and/or remediation of toxic inorganic contaminants such as chromate (CrO42−), perchlorate (ClO4−), nitrate (NO3−), and arsenate (AsO43−) in water, brine, and soil.
Chromium has been widely detected in groundwater and soils, particularly at sites associated with metal plating, wood processing, leather tanning, metal corrosion inhibition, and pigment production. From 1987 to 1993, releases of various chromium compounds to land and water in the U.S. totaled nearly 200 million pounds (EPA, 2006). Compared to the much less soluble Cr(III) species, Cr(VI) species is much more mobile, toxic and carcinogenic. To reduce human exposure to chromium, the U.S. environmental Protection Agency (EPA) has set a maximum contaminant level (MCL) of 0.1 mg/L for total chromium in drinking water.
Traditionally, Cr(VI) is removed from water through reduction of Cr(VI) to Cr(III) using a reducing agent such as ferrous sulfate, sulfur dioxide, or sodium bisulfite, followed by precipitation as Cr(III). In recent years, researchers also demonstrated that Cr(VI) can be effectively reduced by Fe(II) according to the following generic reaction scheme:Cr(VI)+3Fe(II)→Cr(III)+3Fe(III)  (1)
Reduction of Cr(VI) to Cr(III) by powder or granular zero valent iron (ZVI) particles and non-stabilized or agglomerated nanoparticles has been investigated in a number of laboratory and field studies. For example, permeable reactive barriers (PRBs) with commercially available ZVI powder have elicited great interest for in situ treatment of groundwater contaminated with various redox active compounds including Cr(VI). A field-scale PRB using granular ZVI particles to remove Cr(VI) from groundwater was installed at the U.S. Coast Guard Support Center in North Carolina in June of 1996. After eight years of operation, the PRB remained effective for reducing more than 1,500 μg/L of Cr(VI) to less than 1 μg/L.
Metal reduction by zero-valent iron particles follows the general pseudo-first order equation (Ponder, et al, 2002):v=kAs[Me]  (2)where v is the reaction rate, k is the rate constant (M−1m2s−1), [Me] is the metal ion concentration (M), and As is the specific surface area of the iron particles (m2/g). Eqn (2) indicates that the reaction rate is directly proportional to the specific surface area of the ZVI particles. Consequently, reducing particle size is expected to greatly enhance the reaction rate exponentially. For example, reducing the particle size from 10 μm to 10 nm can potentially increase the specific surface area, and thus, the reaction rate, by six orders of magnitude. Ponder et al. (2000) tested a class of resin-supported ZVI nanoparticles (Ferragels, 10-30 nm in diameter) to reduce Cr(VI) in aqueous solutions, and they observed that the reduction of Cr(VI) is 20-30 times faster than the commercial iron filings or iron powder per unit mass of Fe applied. Cao and Zhang (2006) tested non-stabilized ZVI nanoparticles for reduction and immobilization of Cr(VI) in ore processing samples, and observed that the surface-area-normalized reaction rate constant of Cr(VI) reduction by the non-stabilized nanoparticles was about 25 times greater than that by iron powders (100 mesh).
However, ZVI nanoparticles prepared using traditional methods tend to either agglomerate rapidly or react quickly with the surrounding media (e.g. dissolved oxygen or water), resulting in rapid loss in soil mobility as well as reactivity. Because agglomerated ZVI particles are often in the range of micron scale, they are essentially not transportable or deliverable in soils, and thus, cannot be used for in situ applications.
To control nanoparticle agglomeration, various particle stabilizing strategies have been reported. Chen et al. (2004) prepared ZVI nanoparticles with cetylpyridinium chloride (CPC) as a stabilizer for nitrate removal from water. The nanoparticles were reported to have a specific surface area of 25.4 m2/g. He and Zhao (2005b) prepared a new class of starch-stabilized bimetallic nanoparticles to degrade TCE and PCBs. The starched nanoparticles offered a surface area of about 55 m2/g (He and Zhao, 2005b). Schrick et al. (2004; 2002) observed that addition of hydrophilic carbon or poly(acrylic acid) as a supporting agent can enhance the permeability of ZVI nanoparticles in sand and soils.
More recently, He and Zhao (2005a and 2005b) and He et al. (2006) developed a technique for preparing stabilized palladized iron (Fe—Pd) nanoparticles by applying low concentrations of a starch or carboxymethyl cellulose (CMC) as a stabilizer. The stabilized nanoparticles exhibited marked soil mobility and greater reactivity when used for dechlorination of TCE or PCBs in water.
Perchlorate (ClO4−) has been primarily used in solid rocket fuels. It is also widely used in firework powder, roadside flares, airbag inflators, and fertilizers imported from Chile. Past massive application of perchlorate has left a contamination legacy. Perchlorate has been detected at about 400 sites in groundwater, surface water, soil or public drinking water in more than 35 states across the United States with concentrations ranging from 4 μg·L−1 to more than 3.7 million μg·L−1. Perchlorate has been also detected in milk and bottled water.
When ingested, perchlorate can alter the endocrine function by blocking iodide from entering a person's thyroid gland, thereby reducing the production of thyroid hormones. The adverse health effects are reported to be more profound for newborns, children, and pregnant women. A study from the U.S. Centers for Disease Control (CDC) found that of the 36 percent of U.S. women with low iodine intake, almost any amount of perchlorate exposure was linked to a significant change in levels of thyroid hormones. To reduce the related human exposure, the US EPA adopted a Drinking Water Equivalent Level (DWEL) of 24.5 μg·L−1. Meanwhile, California adopted a public health goal of 6 μg·L−1 and Massachusetts set the nation's first drinking water standard of 2 μg·L−1 for perchlorate.
Perchlorate is highly water-soluble, non-complexing, non-volatile, and chemically stable. For its unique chemistry, it has been highly challenging to remove perchlorate from water by traditional water treatment approaches. In recent years, various treatment technologies have been developed and/or tested, including biological reduction (Logan et al., 2002; Min et al., 2004; Xu et al., 2003), ion-exchange (IX) (Gu et al., 2001; 2003; Tripp and Clifford, 2004; 2006; Xiong et al., 2006), tailored activated carbon sorption (Chen et al., 2005; Parette et al., 2005), filtration (Yoon et al., 2003), and chemical reduction (Moore et al., 2003; Gu et al., 2006). However, these technologies are constrained with some critical technical and economic drawbacks such as slow degradation kinetics and production of large volumes of concentrated process waste residuals. For instance, IX has been considered as one of the best available technologies for perchlorate removal. While many commercial IX resins can offer high perchlorate sorption capacity, regeneration efficiency of the IX resins has been found prohibitively poor. As a result, current IX processes are often used on a disposable basis (i.e. the resin is disposed of after only one service run) or, when resin regeneration is practiced, it will result in large volumes of spent regenerant brine. The spent regenerant is often characterized with high salinity (e.g. 6% or higher (w/w) NaCl), concentrated perchlorate, and/or mixture of concentrated acids and salts (1 M FeCl3 and 4 M HCl). Because of the highly stressful conditions, biological treatment of the spent brine is rather challenging and very limited. Consequently, cost-effective technologies that can destroy perchlorate in fresh water and regenerant brine are in dire need.
In recent years, zero-valent iron (ZVI), has attracted increasing interest for abiotic dechlorination of chlorinated organic compounds (Wang and Zhang, 1997), removal of nitrate (Huang and Zhang, 2004; Yang and Lee, 2005), chromate (Wilkin et al., 2005), and arsenic (Kanel et al., 2006). Reduction of perchlorate using ZVI has also been explored by a number of researchers. For instance, Moore et al. (2003; 2005) investigated perchlorate removal by commercial iron filings (size=20-100 mesh, surface area=0.08-5.65 m2·g−1). Up to 66% of perchlorate (0.1 mM) was removed in 336 h at a high iron dosage of 1.25 g·mL−1. Gurol and Kim (2000) reported that UV light can accelerate the perchlorate reduction rate with a dose of 20 g·L−1 or higher metallic iron (size=100 mesh, surface area=0.74 m2·g−1). Oh et al. (2006a) reported that 98% of perchlorate in water was reduced by cast iron (surface area=1.29 m2·g−1) in 1 h at 200° C. via microwave heating and at pH 7.4. In another study, Oh et al. (2006b) reported that perchlorate in wastewater was completely removed by iron at an elevated temperature of 150° C. in 6 hours without pH control. Cao et al. (2005) observed that non-stabilized iron particles were able to degrade perchlorate at temperatures from 25 to 75° C. and determined an activation energy of 79.02±7.75 kJ·mol−1 for the reaction.
Because ZVI-based reactions are surface-mediated processes, increasing surface area of Fe(0) was found to increase the reaction rates (Wang and Zhang, 1997). It has been reported that decreasing the size of Fe(0) particles to the nanoscale can greatly enhance the reaction rates for perchlorate reduction (Cao et al., 2005) and nitrate removal (Yang and Lee, 2005). However, because nanoscale ZVI particles tend to agglomerate rapidly (in a few minutes), “nanoparticles” without a stabilizer are actually agglomerates or flocs of ZVI particles in the scale of micron or sub-millimeter.
To maximize the reactivity of Fe(0) nanoparticles, He and Zhao (2005) and He et al. (2007) modified the conventional water-based approach for preparing ZVI nanoparticles by applying a food-grade water-soluble starch or sodium carboxymethyl cellulose (CMC) as a stabilizer. Compared to conventional non-stabilized ZVI “nanoparticles”, the stabilized nanoparticles displayed much greater surface area, superior physical stability and much faster TCE-dechlorination rates.
Nitrate contamination of groundwater is also a widespread environmental problem, and has been associated with agricultural land runoff, leaching of nitrogen fertilizers, concentrated animal feeding operations, food processing, and industrial waste effluent discharge. Each year, about 11.5 million tons of nitrogen is applied as fertilizer in agricultural areas of the United States. Commercial fertilizer uses in the United States increased by a factor of 20 between 1945 and 1985. Manure produced yearly by farm animals in the United States contributes an estimated 6.5 million tons of nitrogen. The National Research Council (NRC, 1994) reported that there were approximately 300-400 thousands of nitrate-contaminated sites in the United States.
Ingestion of nitrate in drinking water by infants can cause dangerously low oxygen levels in the blood. Nitrate-N concentrations of 4 mg/L or more in rural drinking water supplies have been associated with increased risk of non-Hodgkin's lymphoma. The US EPA has established a maximum contaminant level (MCL) of 10 mg/L nitrate as N (US EPA, 1995). Although nitrate concentration in natural groundwater is generally less than 2 mg/L, it is common for ground water in 10%-25% of the water-supply wells in large regions of the U.S. to exceed the MCL for nitrate.
Because of our tremendous dependence of groundwater, cost-effective remediation of nitrate-contaminated groundwater has been consistently sought for decades. Among the most cited technologies for nitrate removal are ion exchange (IX), biological denitrification, membrane process, and chemical reduction. Although IX-selective resin has been commercially available and IX is an EPA-designated best available technology, IX does not degrade nitrate but rather concentrates nitrate in spent regenerant brine, which demands further costly handling and treatment. Membrane process is another commonly used separation/concentration process. In addition to its prohibitive process cost, disposal of nitrate-laden membrane rejects remains to be a costly obstacle. Biological denitrification has been a rather mature technology for nitrate removal from municipal wastewater. However, it has not gained popularity in drinking water treatment for its slow kinetics under typical drinking water conditions, pH sensitivity, and unfavorable byproducts including taste and odor in the treated water. Thermal nitrate destruction requires an anaerobic condition at temperatures of 200-350° C. and pressures of 600-2800 psig and at pH 13 with the presence of reducing agents, such as ammonia, formate, urea, glucose, methane, and hydrogen. In recent years, a number of studies on reduction of nitrate by zero-valent-iron (ZVI) have been reported (Alowitz and Scherer, 2002; Choe et al., 2000; Huang and Zhang, 2004; Mishra and Farrell, 2005). According to these studies, nitrate is reduced to nitrite, nitrogen gas, and/or ammonia by ZVI following the one of more of the reactions below:Fe0+NO3−+2H+→NO2−+Fe2++H2O  (3)NO2−3Fe0+8H+→3Fe2+NH4++2H2O  (4)4Fe0+NO3−+10H+→NH4++4Fe2++3H2O  (5)5Fe0+2NO3−+12H+→N2(g)+5Fe2++6H2O  (6)
Eqn (3) was proposed by Alowitz and Scherer (2002) using iron powers (18-35 mesh) and Fisher Scientific iron fillings (40 mesh) under controlled solution pH (5.5-9.0). Eqn (4) was proposed by Huang et al. (1998) by assuming nitrite is one of the intermediate products of nitrate reduction. Eqn (5) was suggested to be the main nitrate reduction pathway when nanoscale zero-valent iron (BET specific surface area to mass ratio equal to 31.4 m2/g) was used in anaerobic system under ambient conditions with no pH control (Choe et al., 2000).
Studies have demonstrated that solution pH plays an important role in nitrate reduction by Fe0. Nitrate reduction by iron powder at near-neutral pH was negligible in an unbuffered system, but it was greatly enhanced in the presence of a pH buffer (Zhang and Huang, 2005; Cheng et al., 1997). Ruangchainikom et al. (2006) used a CO2-bubbled system to create an acidic environment favorable to nitrate reduction and found that the bubbling of CO2 flow rate at 200 mL/min was sufficient for supplying H+ to these reactions.
Several strategies have been used to speed up nitrate reduction by ZVI including 1) iron surface pretreatment; 2) addition of selected cations; 3) deposition of a second metal on iron surface; 4) presence of ultraviolet (UV) light and hydrogen peroxide (H2O2); and 5) increasing ZVI surface area by preparing nanoscale ZVI. Liou et al., (2005a) pretreated iron powder (99.6%, electrolytic and finer than 100 mesh) surface with a flow of H2/N2 (20 vol %, 50 mL/min) and maintained at 400° C. for 3 hours to remove the surface passive oxide layers. With this pretreatment, nitrate reduction rate was doubled. Huang and Zhang (2005) found that adding certain selected cations (Fe2+, Fe3+, or Al3+) in feed solution could significantly enhance nitrate reduction. A second metal, such as copper, was loaded onto iron surface as a catalyst for nitrate reduction. However, considerable amounts of nitrite were released (Liou et al., 2005a; 2005b).
Another effective strategy to enhance nitrate reduction by ZVI is to reduce the ZVI particle size, thereby increasing the particle surface area and reactivity. However, current iron nanoparticles, which are typically prepared following the classical borohydride reduction of ferrous or ferric ions in water, tend to agglomerate to large flocs (micrometer to millimeter scale) and precipitate in minutes. Because of the agglomeration, the unique advantage (e.g. high surface area and high reactivity) of nanoscale iron particles is diminished. To prevent the agglomeration, Chen et al. (2005) applied polyvinyl pyrolidine (PVP) and a cationic surfactant cetylpyridinium chloride (CPC) as stabilizers. When used for nitrate reduction, the stabilized iron particles were able to remove 60%-78% nitrate (20 mg/L as NO3−—N) within 10 hours at a iron dosage of 0.5 g/L (12.7 m2/L) under pH 4-7. Recently, He and Zhao developed a new class of Fe—Pd bimetallic nanoparticles by modifying the conventional preparation approach by using an environment-friendly and low-cost starch and a food-grade cellulose (known as sodium carboxymethyl cellulose, NaCMC) as a stabilizer. The stabilized nanoparticles displayed both superior physical stability and much faster reactivity than their non-stabilized counterparts when used for degradation of chlorinated hydrocarbons (He and Zhao, 2005; 2006).
Compared to nitrate removal from fresh water, research on nitrate reduction in saline water has been very limited and remains in its exploratory stage. Biological denitrification has been found effective to denitrify nitrate in seawater (Labelle et al., 2005) and in ion exchange regenerant brine containing 1%-12.5% NaCl (Clifford and Liu, 1993b; Okeke et al., 2002; Peyton et al., 2001). However, Clifford and Liu (1993b) reported a 10% drop in denitrification rate in 0.5 N NaCl than in fresh-water controls a bench-scale biological reactor was used to treat ion-exchange brine and time for >95% denitrification was 8 hours. Earlier, Van der Hoek et al. (1987) reported a combined ion exchange/biological denitrification process for nitrate removal from ground water, in which nitrate was removed by ion exchange and the nitrate-laden regenerant brine was denitrified by a biological denitrification reactor. But a decrease in denitrification capacity was observed when high NaCl concentration (10-30 g/L) presented. Bench-scale sequencing batch reactors using activated sludge have been reported to remove a wastewater containing 36,000 mg/L NO3− with ionic strength of 3.0 (18% total dissolved solids) and both nitrite and nitrate reduction rates reduced with increasing salinity (Glass and Silverstein, 1999). Peyton et al. (2001) reported specific nitrate reduction rate coefficients in a range from 1.20×10−2±7.22×10−4 (L/h mg TSS) to 5.54×10−3±3.94×10−4 (L/h mg TSS) depending on carbon sources in a pH 9 solution containing 12.5% NaCl. To date, there has been no research reported on the abiotic degradation of nitrate in saline water using stabilized ZVI.
Arsenic in soils and groundwater results from natural sources (e.g. natural geochemical reactions) as well as anthropogenic activities, such as mining, discharges of industrial wastes, military activities, and application of agricultural pesticides. Arsenic is ranked the second most common inorganic pollutant in the U.S. superfund sites. Arsenic-contaminated soils, sediments and waste slurry are major sources of arsenic in food and water. To mitigate the toxic effect on human health, the maximum contaminant level (MCL) for arsenic in drinking water was lowered from the previous 50 ppb to 10 ppb, effective in January 2006.
Arsenic is a redox active element, with As(V) or (III) being the two most common stable oxidation states in soils. In general, inorganic arsenic is more toxic than organic arsenic, and arsenic in soils is less bioavailable and less bioaccessible than As in water due to soil adsorption effect.
Arsenate can strongly interact with soils, especially, iron (hydr)oxides. Adsorption of arsenate by iron (hydr)oxides have been widely studied. These studies have focused on the adsorption and surface complexation of arsenic on the amorphous and crystalline iron oxide structures, such as ferrihydrite and goethite. The complexation between arsenate and iron (hydr)oxide surfaces has been known to be inner-sphere surface complexation as either monodentate sharing, bidentate sharing, or bidentate edge sharing complexes.
Laboratory-scale and field-scale studies have been reported on in situ remediation of As-contaminated groundwater by zero-valent iron (ZVI) (Nikolaidis et al. 2003; Su and Plus 2001; Su and Puls 2001) and iron oxides (Fendorf et al. 1997). They observed that ZVI can reduce the concentration of As in aqueous phase. Recently, nanoscale iron-based media (such as zero-valent iron) have been studied for potential uses in environmental remediation (Huber 2005; Jegadeesan et al. 2005; Zhang 2003). Because of the small particle size, large surface area, and high reactivity, these nanoscale materials have showed great potential for treatment of contaminated soil and groundwater (Chen et al. 2005; Joo et al. 2004; Yang and Lee 2005). Cumbal and Sengupta (Cumbal and Sengupta 2005) studied arsenic removal from water by hydrated iron oxides nanoparticles loaded on polymer-matrix, and the immobilized nanoscale iron oxides displayed high sorption capacity for both arsenite and arsenate. For arsenic removal in groundwater by iron-based nanoparticles, surface adsorption appears to be an important mechanism (Kanel et al. 2006). Compared to commercial iron powder or granular iron particles, ZVI nanoparticles offer much faster sorption kinetics and are more deliverable in the subsurface. Consequently, iron nanoparticles hold great potential to immobilize arsenic in situ in contaminated soil and groundwater.
However, as previously noted herein, because of the high reactivity and inter-particle interactions, ZVI nanoparticles tend to agglomerate rapidly, resulting in the formation of much larger aggregated particles and loss of reactivity and soil mobility. To prevent iron nanoparticle agglomeration, various particle stabilization strategies were reported (He and Zhao 2005; Ponder et al. 2001; Raveendran et al. 2003). He and Zhao (2005, 2006) reported a new method for synthesizing stabilized iron nanoparticles by using some low-cost and environmentally benign starch and cellulose as a stabilizer. The stabilized nanoparticles displayed much improved physical stability, soil mobility, and reactivity compared to non-stabilized iron particles.
To quantify relative As mobility and leachability in soil, two operationally defined measures, bioaccessibility and TCLP (toxicity characteristic leaching procedure) leachability, have been commonly used. Bioaccessibility is quantified by a physiologically based extraction test (PBET), which mimics the conditions in human stomach and essentially reflects an in vivo accessibility of As (Ruby et al. 1999). TCLP is an EPA-defined standard method for measuring extractability of various chemicals from solid wastes. Earlier, a number of researchers (Akhter et al. 2000; Jing et al. 2005; Miller et al. 2000) used TCLP tests to evaluate the leachability of As in contaminated soils.
Akhter et al. (2000) concluded that higher iron content in soil reduces the leachability of arsenic. Yang et al. (2002) observed that high iron content reduced the bioaccessibility of arsenic in soil.